necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.
Concise International Chemical Assessment Document 44
SILVER AND SILVER COMPOUNDS:
ENVIRONMENTAL ASPECTS
The layout and pagination of this pdf file are not necessarily identical to those of the hard copy
First draft prepared by Mr P.D. Howe and Dr S. Dobson, Centre for Ecology and Hydrology, Monks Wood, United Kingdom
Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of
Chemicals.
World Health Organization
Geneva, 2002
Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.
The
Inter-Organization Programme for the Sound Management of Chemicals (IOMC) wasestablished in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.
WHO Library Cataloguing-in-Publication Data Silver and silver compounds: environmental aspects.
(Concise international chemical assessment document ; 44) 1.Silver - adverse effects 2.Water pollutants, Chemical
3.Risk assessment 4.Environmental exposure I.International Programme on Chemical Safety II.Series
ISBN 92 4 153044 8 (NLM Classification: QV 297) ISSN 1020-6167
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The Federal Ministry for the Environment, Nature Conservation and Nuclear Safety, Germany, provided financial support for the printing of this publication.
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iii
FOREWORD...1
1. EXECUTIVE SUMMARY...4
2. IDENTITY AND PHYSICAL/CHEMICAL PROPERTIES ...5
3. ANALYTICAL METHODS...5
4. SOURCES OF ENVIRONMENTAL EXPOSURE...6
5. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION...7
6. ENVIRONMENTAL LEVELS...11
7. EFFECTS ON ORGANISMS IN THE LABORATORY AND FIELD ...13
7.1 Aquatic environment ...13
7.2 Terrestrial environment...18
8. EFFECTS EVALUATION ...19
9. PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES...21
REFERENCES...22
APPENDIX 1 — SOURCE DOCUMENT...27
APPENDIX 2 — CICAD PEER REVIEW ...27
APPENDIX 3 — CICAD FINAL REVIEW BOARD...28
INTERNATIONAL CHEMICAL SAFETY CARDS...29
RÉSUMÉ D’ORIENTATION...33
RESUMEN DE ORIENTACIÓN ...35
FOREWORD
Concise International Chemical Assessment Documents (CICADs) are the latest in a family of publications from the International Programme on Chemical Safety (IPCS) — a cooperative programme of the World Health Organization (WHO), the International Labour Organization (ILO), and the United Nations Environment Programme (UNEP). CICADs join the Environmental Health Criteria documents (EHCs) as authoritative documents on the risk assessment of chemicals.
International Chemical Safety Cards on the relevant chemical(s) are attached at the end of the CICAD, to provide the reader with concise information on the protection of human health and on emergency action. They are produced in a separate peer-reviewed procedure at IPCS.
CICADs are concise documents that provide sum- maries of the relevant scientific information concerning the potential effects of chemicals upon human health and/or the environment. They are based on selected national or regional evaluation documents or on existing EHCs. Before acceptance for publication as CICADs by IPCS, these documents undergo extensive peer review by internationally selected experts to ensure their completeness, accuracy in the way in which the original data are represented, and the validity of the conclusions drawn.
The primary objective of CICADs is characteri- zation of hazard and dose–response from exposure to a chemical. CICADs are not a summary of all available data on a particular chemical; rather, they include only that information considered critical for characterization of the risk posed by the chemical. The critical studies are, however, presented in sufficient detail to support the conclusions drawn. For additional information, the reader should consult the identified source documents upon which the CICAD has been based.
Risks to human health and the environment will vary considerably depending upon the type and extent of exposure. Responsible authorities are strongly encour- aged to characterize risk on the basis of locally measured or predicted exposure scenarios. To assist the reader, examples of exposure estimation and risk characteriza- tion are provided in CICADs, whenever possible. These examples cannot be considered as representing all possible exposure situations, but are provided as guidance only. The reader is referred to EHC 170.1
1International Programme on Chemical Safety (1994) Assessing human health risks of chemicals: derivation of guidance values for health-based exposure limits. Geneva, World Health Organization (Environmental Health Criteria 170) (also available at http://www.who.int/pcs/).
While every effort is made to ensure that CICADs represent the current status of knowledge, new informa- tion is being developed constantly. Unless otherwise stated, CICADs are based on a search of the scientific literature to the date shown in the executive summary. In the event that a reader becomes aware of new informa- tion that would change the conclusions drawn in a CICAD, the reader is requested to contact IPCS to inform it of the new information.
Procedures
The flow chart on page 2 shows the procedures followed to produce a CICAD. These procedures are designed to take advantage of the expertise that exists around the world — expertise that is required to produce the high-quality evaluations of toxicological, exposure, and other data that are necessary for assessing risks to human health and/or the environment. The IPCS Risk Assessment Steering Group advises the Coordinator, IPCS, on the selection of chemicals for an IPCS risk assessment based on the following criteria:
• there is the probability of exposure; and/or
• there is significant toxicity/ecotoxicity.
Thus, a priority chemical typically
• is of transboundary concern;
• is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management;
• is significantly traded internationally;
• has high production volume;
• has dispersive use.
The Steering Group will also advise IPCS on the appropriate form of the document (i.e., EHC or CICAD) and which institution bears the responsibility of the document production, as well as on the type and extent of the international peer review.
The first draft is based on an existing national, regional, or international review. Authors of the first draft are usually, but not necessarily, from the institution that developed the original review. A standard outline has been developed to encourage consistency in form.
The first draft undergoes primary review by IPCS to ensure that it meets the specified criteria for CICADs.
The second stage involves international peer review by scientists known for their particular expertise and by scientists selected from an international roster compiled by IPCS through recommendations from IPCS national Contact Points and from IPCS Participating Institutions.
Adequate time is allowed for the selected experts to undertake a thorough review. Authors are required to take reviewers’ comments into account and revise their
CICAD PREPARATION FLOW CHART
Selection of priority chemical, author institution, and agreement
on CICAD format
9
Preparation of first draft
9
Primary acceptance review by IPCS and revisions as necessary
9
Selection of review process
9
Peer review
9
Review of the comments and revision of the
document
9
Final Review Board:
Verification of revisions due to peer review comments, revision, and approval of the document
9
Editing
Approval by Coordinator, IPCS
9
Publication of CICAD on Web and as printed text
Advice from Risk Assessment Steering Group
Criteria of priority:
$ there is the probability of exposure;
and/or
$ there is significant toxicity/ecotoxicity.
Thus, it is typical of a priority chemical that
$ it is of transboundary concern;
$ it is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management;
$ there is significant international trade;
$ the production volume is high;
$ the use is dispersive.
Special emphasis is placed on avoiding duplication of effort by WHO and other international organizations.
A prerequisite of the production of a CICAD is the availability of a recent high-quality national/regional risk assessment document = source document. The source document and the CICAD may be produced in parallel. If the source document does not contain an environ- mental section, this may be produced de novo, provided it is not controversial. If no source document is available, IPCS may produce a de novo risk assessment document if the cost is justified.
Depending on the complexity and extent of controversy of the issues involved, the steering group may advise on different levels of peer review:
$ standard IPCS Contact Points
$ above + specialized experts
$ above + consultative group
draft, if necessary. The resulting second draft is submitted to a Final Review Board together with the reviewers’ comments. At any stage in the international review process, a consultative group may be necessary to address specific areas of the science.
The CICAD Final Review Board has several important functions:
• to ensure that each CICAD has been subjected to an appropriate and thorough peer review;
• to verify that the peer reviewers’ comments have been addressed appropriately;
• to provide guidance to those responsible for the preparation of CICADs on how to resolve any remaining issues if, in the opinion of the Board, the author has not adequately addressed all comments of the reviewers; and
• to approve CICADs as international assessments.
Board members serve in their personal capacity, not as representatives of any organization, government, or industry. They are selected because of their expertise in human and environmental toxicology or because of their experience in the regulation of chemicals. Boards are chosen according to the range of expertise required for a meeting and the need for balanced geographic repre- sentation.
Board members, authors, reviewers, consultants, and advisers who participate in the preparation of a CICAD are required to declare any real or potential conflict of interest in relation to the subjects under discussion at any stage of the process. Representatives of nongovernmental organizations may be invited to observe the proceedings of the Final Review Board.
Observers may participate in Board discussions only at the invitation of the Chairperson, and they may not participate in the final decision-making process.
1. EXECUTIVE SUMMARY
This CICAD on silver and silver compounds (environmental aspects) was prepared by the Centre for Ecology and Hydrology, Monks Wood, United
Kingdom. It is based on the US Department of the Interior’s Contaminant Hazard Reviews report entitled Silver hazards to fish, wildlife, and invertebrates: A synoptic review (Eisler, 1997), updated with reference to Eisler (2000) and supplemented by a literature search (April 2001). This document does not cover human health effects of silver; a review of human health aspects can be found in ATSDR (1990). Information on the nature of the peer review and availability of the source document is presented in Appendix 1. Information on the peer review of this CICAD is presented in Appendix 2. This CICAD was approved as an international assess- ment at a meeting of the Final Review Board, held in Ottawa, Canada, on 29 October – 1 November 2001.
Participants at the Final Review Board meeting are listed in Appendix 3. The International Chemical Safety Cards for silver (ICSC 0810) and silver nitrate (ICSC 1116), produced by the International Programme on Chemical Safety (IPCS, 1999a, 1999b), have also been reproduced in this document.
Silver is a rare but naturally occurring metal, often found deposited as a mineral ore in association with other elements. Emissions from smelting operations, manufacture and disposal of certain photographic and electrical supplies, coal combustion, and cloud seeding are some of the anthropogenic sources of silver in the biosphere. The global biogeochemical movements of silver are characterized by releases to the atmosphere, water, and land by natural and anthropogenic sources, long-range transport of fine particles in the atmosphere, wet and dry deposition, and sorption to soils and sedi- ments.
The most recent measurements of silver in rivers, lakes, and estuaries using clean techniques show levels of about 0.01 µg/litre for pristine, unpolluted areas and 0.01–0.1 µg/litre in urban and industrialized areas. Silver concentrations reported prior to the implementation of ultra-clean metal sampling, which began in the late 1980s, should be treated with caution. Maximum con- centrations of total silver recorded during the 1970s and 1980s in selected non-biological materials were 36.5 ng/m3 in air near a smelter; 2.0 µg/m3 in atmos- pheric dust; 0.1 µg/litre in oil well brines; 4.5 µg/litre in precipitation from clouds seeded with silver iodide;
6.0 µg/litre in groundwater near a hazardous waste site;
8.9 µg/litre in seawater from Galveston Bay, USA;
260 µg/litre near photographic manufacturing waste discharges; 300 µg/litre in steam wells; 300 µg/litre in treated photoprocessing wastewaters; 31 mg/kg in soils;
43 mg/litre in water from certain hot springs; 50 mg/kg
in granite; as much as 100 mg/kg in crude oils; and 150 mg/kg in river sediments. It should be noted that levels of silver in the environment have declined; for example, in the Lower Genesee River, USA, near a photographic manufacturing plant, levels declined from 260 µg/litre in the 1970s to below the detection limit (<10 µg/litre) in the 1990s. It should also be noted that only a small portion of the total silver in each of the environmental compartments is biologically available.
The ability to accumulate dissolved silver varies widely between species. Some reported bioconcentration factors for marine organisms (calculated as milligrams of silver per kilogram fresh weight organism divided by milligrams of silver per litre of medium) are 210 in diatoms, 240 in brown algae, 330 in mussels, 2300 in scallops, and 18 700 in oysters, whereas bioconcentra- tion factors for freshwater organisms have been reported to range from negligible in bluegills (Lepomis macro- chirus) to 60 in daphnids; these values represent uptake of bioavailable silver in laboratory experiments. Labora- tory studies with the less toxic silver compounds, such as silver sulfide and silver chloride, reveal that accumula- tion of silver does not necessarily lead to adverse effects.
At concentrations normally encountered in the environ- ment, food-chain biomagnification of silver in aquatic systems is unlikely. Elevated silver concentrations in biota occur in the vicinities of sewage outfalls, electro- plating plants, mine waste sites, and silver iodide-seeded areas. Maximum concentrations recorded in field collec- tions, in milligrams total silver per kilogram dry weight (tissue), were 1.5 in marine mammals (liver) (except Alaskan beluga whales Delphinapterus leucas, which had concentrations 2 orders of magnitude higher than those of other marine mammals), 6 in fish (bone), 14 in plants (whole), 30 in annelid worms (whole), 44 in birds (liver), 110 in mushrooms (whole), 185 in bivalve molluscs (soft parts), and 320 in gastropods (whole).
In general, silver ion was less toxic to freshwater aquatic organisms under conditions of low dissolved silver ion concentration and increasing water pH, hard- ness, sulfides, and dissolved and particulate organic loadings; under static test conditions, compared with flow-through regimens; and when animals were ade- quately nourished instead of being starved. Silver ions are very toxic to microorganisms. However, there is generally no strong inhibitory effect on microbial activity in sewage treatment plants because of reduced bioavailability due to rapid complexation and adsorp- tion. Free silver ion was lethal to representative species of sensitive aquatic plants, invertebrates, and teleosts at nominal water concentrations of 1–5 µg/litre. Adverse effects occur on development of trout at concentrations as low as 0.17 µg/litre and on phytoplankton species composition and succession at 0.3–0.6 µg/litre.
A knowledge of the speciation of silver and its consequent bioavailability is crucial to understanding the potential risk of the metal. Measurement of free ionic silver is the only direct method that can be used to assess the likely effects of the metal on organisms. Speciation models can be used to assess the likely proportion of the total silver measured that is bioavailable to organisms.
Unlike some other metals, background freshwater concentrations in pristine and most urban areas are well below concentrations causing toxic effects. Levels in most industrialized areas border on the effect concen- tration, assuming that conditions favour bioavailability.
On the basis of available toxicity test results, it is unlikely that bioavailable free silver ions would ever be at sufficiently high concentrations to cause toxicity in marine environments.
In general, accumu lation of silver by terrestrial plants from soils is low, even if the soil is amended with silver-containing sewage sludge or the plants are grown on tailings from silver mines, where silver accumulates mainly in the root systems. No data were found on effects of silver on wild birds or mammals. Silver was harmful to poultry (tested as silver nitrate) at concen- trations as low as 100 mg total silver/litre in drinking- water or 200 mg total silver/kg in diets. Sensitive laboratory mammals were adversely affected at total silver concentrations (added as silver nitrate) as low as 250 µg/litre in drinking-water (brain histopathology), 6 mg/kg in diet (high accumulations in kidneys and liver), or 13.9 mg/kg body weight (lethality).
2. IDENTITY AND PHYSICAL/CHEMICAL PROPERTIES
Silver is a white, ductile metal occurring naturally in the pure form and in ores. Some silver compounds are extremely photosensitive and are stable in air and water, except for tarnishing readily when exposed to sulfur compounds. Metallic silver is insoluble in water, but many silver salts, such as silver nitrate (AgNO3), are soluble. In the natural environment, silver occurs pri- marily in the form of the sulfide (Ag2S) or is intimately associated with other metal sulfides, especially those of lead, copper, iron, and gold, which are all essentially insoluble (ATSDR, 1990). The colloidal phase, com- posed of organic and inorganic materials, is the most important component of aqueous silver (Bell & Kramer, 1999). Monovalent silver ion (Ag+) is rare or negligible in the natural environment. Silver readily forms com- pounds with antimony, arsenic, selenium, and tellurium (Smith & Carson, 1977). Silver has two stable isotopes (107Ag and 109Ag) and 20 radioisotopes; none of the radioisotopes of silver occurs naturally. Several com- pounds of silver are potential explosion hazards: silver
oxalate (Ag2C2O4) decomposes explosively when heated; silver acetylide (Ag2C2) is sensitive to detonation on contact; and silver azide (AgN3) detonates sponta- neously under certain circumstances (Smith & Carson, 1977).
Selected physicochemical properties of silver and its salts are summarized in Table 1. Additional properties for silver (ICSC 0810) and silver nitrate (ICSC 1116) are presented in the International Chemical Safety Cards reproduced in this document.
3. ANALYTICAL METHODS
A variety of spectrographic, colorimetric, polaro- graphic, and other analytical techniques are used for routine measurement of silver in biological and abiotic samples. Silver concentrations reported prior to the implementation of ultra-clean metal sampling, which began in the late 1980s, should be treated with caution.
Atomic absorption and plasma emission spectros- copy are perhaps the most widely used analytical techniques for the determination of silver levels in air, soil, and water. Rains et al. (1984) employed atomic absorption spectroscopy with flame atomization and direct current plasma–atomic emission spectroscopy to determine silver levels in solid waste leachate. Detection limits of silver in leachate samples by the two techniques were 0.473 µg/ml and 0.38 µg/litre, respectively.
Inductively coupled argon plasma with atomic emission spectroscopy has been recommended for determining silver in air and for analysing dissolved, suspended, or total silver in drinking-water, surface water, and domestic and industrial wastewaters.
Detection limits were 26 ng/ml in air and 7.0 µg/litre in water (ATSDR, 1990). Inductively coupled plasma–
mass spectrometry is used to measure silver in environ- mental media at a detection limit of 0.4 ng/litre.
Sensitive voltammetric techniques using anodic stripping have been developed to measure free silver ion in solution at concentrations as low as 0.1 µg/litre (Schildkraut, 1993; Song & Osteryoung, 1993; Schild- kraut et al., 1998). However, the anodic stripping voltammetric method does not work well with natural samples containing large amounts of organic matter, such as those found in sewage treatment plants (Ownby et al., 1997).
Trace levels of silver (10–6 to 10–9 g/g of sample) can be accurately determined in biological samples by several different analytical techniques. These methods include high-frequency plasma torch–atomic emission
Table 1: Physicochemical properties of silver and some of its salts.
Properties Silver Silver nitrate Silver sulfide Silver chloride
Alternative names Argentum; argentum crede Cl 77820; shell silver; silver atom; silver colloidal; silflake;
silpowder; silber
Lunar caustic fused silver nitrate; moulded silver nitrate argenti; nitras; nitric acid silver (I) salt; nitric acid silver (1+) salt; silver (1+) nitrate
Acanthite; argentous sulfide
Silver (I) chloride;
silver monochloride
Chemical Abstracts Service number
7440-22-4 7761-88-8 21548-73-2 7783-90-6
Chemical formula Ag AgNO3 Ag2S AgCl
Molecular mass 107.87 169.89 247.80 143.34
Physical state Solid metal Solid crystalline Grey-black solid White solid
Boiling point 2212 °C Decomposes at 440 °C Decomposes at 810 °C 1550 °C
Solubility in water (at 20 °C)
Insoluble 2160 g/litre Insoluble 1.93 mg/litre
Solubility in organic solvents/acids
Soluble in nitric acid but not sulfuric acid
Soluble in ethanol and acetone
Density (at 20 °C) 10.5 g/cm3 4.35 g/cm3 7.33 g/cm3 5.56 g/cm3
spectroscopy, neutron activation analysis, graphite furnace (flameless) atomic absorption spectroscopy, flame atomic absorption spectroscopy, and micro-cup atomic absorption spectroscopy. Atomic absorption spectroscopy equipped with various atomizers is the most prevalent analytical method used. Graphite furnace atomic absorption spectroscopy offers high detectability (subnanogram per gram sample) and requires relatively small samples for analysis of biological tissues
(DiVincenzo et al., 1985). Detection limts of 0.02 µg/g, 0.2 µg/g, 0.005 µg/litre, and 0.5 µg/100 ml were achieved for hair, faeces, urine, and blood, respectively.
Starkey et al. (1987) modified the graphite furnace atomic absorption spectroscopy technique for deter- mining trace levels of silver in blood samples and reported a detection limit of 15 ng/100 ml.
Analytical techniques to measure trace levels of silver are extremely complex, and, unless the proper sample containers are used, the silver in a solution can be lost to container walls within hours after collection (Wen et al., 1997). Wen et al. (1997) found that ultra- violet irradiation of the sample water consistently achieved 100% recovery. They also found that the ultraviolet irradiation method was required for silver analysis in concentrated colloidal solutions. The presence of organic matter and sulfide clusters has been shown to enhance the loss of silver in natural water.
4. SOURCES OF ENVIRONMENTAL EXPOSURE
Silver is a rare but naturally occurring metal, often found deposited as a mineral ore in association with other elements (ATSDR, 1990). Argentite is the main
ore from which silver is extracted by cyanide, zinc reduction, or electrolytic processes (Fowler & Nordberg, 1986). Silver is frequently recovered as a by-product from smelting of nickel ores in Canada, from lead-zinc and porphyry copper ores in the USA, and from plati- num and gold deposits in South Africa (Smith & Carson, 1977). About 12–14% of the domestic silver output is recovered from lead ores, and about 4% from zinc ores.
Secondary sources of silver include new scrap generated in the manufacture of silver-containing products; coin and bullion; and old scrap from electrical products, old film and photoprocessing wastes, batteries, jewellery, silverware, and bearings (Smith & Carson, 1977).
World production of silver increased from 7.4 mil- lion kilograms in 1964 to 9.1 million kilograms in 1972 and to 9.7 million kilograms in 1982 (Fowler & Nord- berg, 1986). In 1986, 13.06 million kilograms of silver were produced globally; the USA produced 1.1 million kilograms in 1986 but consumed 3.9 million kilograms (ATSDR, 1990). In 1990, the estimated world mine production of silver was 14.6 million kilograms. Major producers were Mexico, with 17% of the total; the USA, with 14%; Peru, with 12%; the former Soviet Union, with 10%; and Canada, with 9% (Eisler, 1997). Ratte (1998) gives a world production figure of 14.9 million kilograms for 1995, whereas the World Silver Survey 2000 states that mine production of silver for 1999 was 15.5 million kilograms (Silver Institute, 2000).
Emissions from smelting operations, manufacture and disposal of certain photographic and electrical supplies, coal combustion, and cloud seeding are some of the anthropogenic sources of silver in the biosphere (Eisler, 1997). Fallout from cloud seeding with silver iodide (AgI) is not always confined to local precipita- tion; silver residuals have been detected several hundred
kilometres downwind of seeding events (Freeman, 1979).
In 1978, the estimated loss of silver to the environ- ment in the USA was 2.5 million kilograms, mostly to terrestrial and aquatic ecosystems; the photographic industry alone accounted for about 47% of all silver discharged into the environment from anthropogenic sources (Smith & Carson, 1977). Purcell & Peters (1998) cited a report by Scow et al. (1981) that stated that the estimated loss of silver to the environment in the USA in 1978 ranged from 2.4 to 2.5 million kilograms. Twenty- nine per cent was released to the aquatic environment, whereas 68% was released to land as solid waste. It was reported that 30% of the total release originated from natural sources and 30% from photographic developing and manufacture. In 1999, the estimated release to the environment in the USA via emissions, discharges, and waste disposal from sites listed in the Toxic Release Inventory were 270 000 kg for silver and 1.7 million kilograms for silver compounds. Releases to land amounted to 90% for silver compounds and 40% for silver, whereas nearly 60% of the silver releases were via off-site waste disposal (TRI, 1999).
Most of the silver lost to the environment enters terrestrial ecosystems, where it is immobilized in the form of minerals, metal, or alloys; agricultural lands may receive as much as 80 000 kg of silver per year from photoprocessing wastes in sewage sludge. An estimated 150 000 kg of silver enter the aquatic environment every year from the photographic industry, mine tailings, and electroplating (Smith & Carson, 1977). The atmosphere receives 300 000 kg of silver each year from a variety of sources.
Silver has been used for ornaments and utensils for almost 5000 years and as a precious metal, a monetary medium, and a basis of wealth for more than 2000 years.
In 1990, about 50% of the refined silver consumed in the USA was used to manufacture photographic and X-ray products; 25% in electrical and electronic prod- ucts; 10% in electroplated ware, sterlingware, and jewel- lery; 5% in brazing alloys; and 10% in other uses (Eisler, 1997).
Because of their bacteriostatic properties, silver compounds are used in filters and other equipment to purify swimming pool water and drinking-water and in the processing of foods, drugs, and beverages (ATSDR, 1990).
Silver nitrate has been used for many years as an eyedrop treatment to prevent ophthalmia neonatorum (ATSDR, 1990). Several silver-containing pharmaceuti- cals have been used topically on skin or mucous mem- branes to assist in healing burn patients and to combat
skin ulcers. Oral medicines containing silver include antismoking lozenges containing silver acetate
(AgC2H3O2), breath mints coated with silver, and silver nitrate solutions for treating gum disease (ATSDR, 1990). The widespread medical use of silver compounds for topical application to mucous membranes and for internal use has become nearly obsolete in the past 50 years because of the fear of argyria and the develop- ment of sulfonamide and antibiotic microbials (Smith &
Carson, 1977).
5. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION
The global biogeochemical movements of silver are characterized by releases to the atmosphere, water, and land by natural and anthropogenic sources, long-range transport of fine particles in the atmosphere, wet and dry deposition, and sorption to soils and sediments (ATSDR, 1990). The chief source of silver contamination of water is silver thiosulfate complexes in photographic develop- ing solutions that photofinishers discard directly to sewers (Smith & Carson, 1977). Secondary waste treat- ment converts most of the silver thiosulfate complex to insoluble silver sulfide and forms some metallic silver (Lytle, 1984). About 95% of the total silver is removed in publicly owned treatment works from inputs contain- ing municipal sewage and commercial photoprocessing effluents, and effluents contain less than 0.07 µg ionic silver/litre; concentrations were independent of the influent silver concentration (Lytle, 1984; Shafer et al., 1998). Silver in sewage treatment plant effluents may be associated with suspended particles or be present as thiosulfate complex, colloidal silver complex, colloidal silver chloride (AgCl), silver sulfide, or soluble organic complexes (Smith & Carson, 1977). Silver on suspended matter and in colloidal forms and insoluble salts ulti- mately settles out in the sediments. At the water treat- ment plant, most of the silver is precipitated after treatment with lime or adsorbed after treatment with alum flocculant. Chlorination converts some silver to silver chloride or to a soluble silver chloride complex (Smith & Carson, 1977). Aerobic biodegradation of a photoprocessing wastewater containing 1.85 mg total silver/litre did not adversely affect the activated sludge process (Pavlostathis & Maeng, 1998). Practically all silver became associated with the sludge solids at 1840 mg silver/kg mixed liquor suspended solids. When fresh sludge and aerobically digested sludge solids were subjected to leaching procedures, the resulting silver concentration was at least 40 times lower than the regulatory limit of 5 mg/litre (Pavlostathis & Maeng, 1998).
Forms of silver in atmospheric emissions are probably silver sulfide, silver sulfate (Ag2SO4), silver carbonate (Ag2CO3), silver halides, and metallic silver (Smith & Carson, 1977). About 50% of the silver released into the atmosphere from industrial operations is transported more than 100 km and is eventually deposited in precipitation (ATSDR, 1990).
Emissions of silver from coal-fired power plants may lead to accumulations in nearby soils (Fowler &
Nordberg, 1986). Silver in soils is largely immobilized by precipitation to insoluble salts and by complexation or adsorption by organic matter, clays, and manganese and iron oxides (Smith & Carson, 1977).
Silver occurs naturally in several oxidation states, usually as Ag0 and Ag+; other possible oxidation states of silver are Ag2+ and Ag3+ (ATSDR, 1990). In this report, “silver ion” refers to Ag+, unless stated otherwise.
In surface fresh water, silver may be found as the mono- valent ion; in combination with sulfide, bicarbonate, or sulfate; as part of more complex ions with chlorides and sulfates; and adsorbed onto particulate matter (ATSDR, 1990). In the aqueous phase, silver at the lowest concen- trations exists either as a simple silver sulfhydrate (AgSH) or as a simple polymer HS-Ag-S-Ag-SH (Bell
& Kramer, 1999). At higher concentrations, colloidal silver sulfide or polysulfide complexes are formed.
Silver ion binds strongly with sulfide ion in inorganic and organic species, resulting in nanogram per litre aqueous dissolved concentrations (Bell & Kramer, 1999). Trace levels of dissolved silver in the presence of ferric sulfide are rapidly adsorbed, and silver remaining in solution remains as silver sulfide; however, silver thiolate complexes can be the dominant dissolved species in highly contaminated waters near urban centres or in waters with high levels of natural organic matter (Adams & Kramer, 1998). The most important and crucial aspect of silver thiolate chemistry is the rapid exchange of silver ion among thiolates, whereby silver ion can transfer onto, or off, particulate materials or the cells of an organism. Silver thiolates also react rapidly with hydrogen sulfide or HS– as ligands to form silver sulfide, although the reverse process is slow (Bell &
Kramer, 1999). The monovalent ion does not hydrolyse appreciably in solution and is considered to be a mild oxidizing agent (Smith & Carson, 1977). Hypervalent silver species, such as Ag2+ and Ag3+, are significantly more effective as oxidizing agents than Ag0 and Ag+ (Kouadio et al., 1990; Sun et al., 1991) but unstable in aqueous environments, especially at water temperatures near 100 °C (Smith & Carson, 1977).
In fresh water and soils, the primary silver com- pounds under oxidizing conditions are bromides, chlorides, and iodides; under reducing conditions, the free metal and silver sulfide predominate (ATSDR, 1990). In river water, one study showed silver present as
the monovalent ion at 53–71% of the total silver, as silver chloride at 28–45%, and as silver chloride ion (AgCl–) at 0.6–2.0% (Whitlow & Rice, 1985). Increas- ing salinity of brackish and marine waters increases concentrations of silver-chloro complexes (AgCl0, AgCl2–, AgCl32–
, AgCl43–
); these chloro complexes retain some silver in dissolved form, and relatively small anthropogenic quantities can substantially enrich the environment (Luoma, 1994). In the open ocean, the principal dissolved form of silver is AgCl2–, but the most bioavailable form may be the neutral monochloro complex silver chloride (Bryan & Langston, 1992).
Sorption is the dominant process that controls the partitioning of silver in water and its movement in soils and sediments (US EPA, 1980; ATSDR, 1990). Silver may leach from soils into groundwater, the leaching rate increasing with decreasing pH and increasing drainage (ATSDR, 1990). Silver adsorbs to manganese dioxide, ferric compounds, and clay minerals, and these com- pounds are involved in silver deposition into sediments;
sorption by manganese dioxide and precipitation with halides reduce the concentration of dissolved silver, resulting in higher concentrations in sediments than in the water column (US EPA, 1980). Under reducing conditions, adsorbed silver in sediments may be released and subsequently reduced to metallic silver, or it may combine with reduced sulfur to form the insoluble silver sulfide (US EPA, 1980). In wastewaters, inorganic ligands were found to be important for silver binding.
These inorganic ligands probably consist of metal sulfides, which are stable for hours and days in oxic waters. When complexed to these strong-affinity ligands, Ag(I) is protected from photoreduction to zero-valent silver (Adams & Kramer, 1999a). Partitioning of silver into size fractions showed that a significant proportion of silver is in the colloidal (30–35%) and dissolved phases (15–20%). Dissolved-phase concentrations were rela- tively constant in treatment plant effluent and receiving waters, suggesting that silver is strongly complexed by ligands that are not significantly affected by aggregation or sorption processes (Adams & Kramer, 1999b). How- ever, depending on the conditions, sediments may be a significant source of silver to the water column. In one study, anoxic sediments containing 1–27 g silver/kg dry weight and 10 mmol of acid volatile sulfide/kg dry weight were resuspended in oxygenated seawater for several hours to days. The seawater in contact with sedi- ment containing 10.8 g silver/kg had 20 µg silver/litre;
seawater in contact with sediments containing 27 g silver/kg had about 2000 µg silver/litre, which seems to be the solubility of silver in seawater (Eisler, 1997).
The availability of free silver in marine environ- ments is strongly controlled by salinity, because of the affinity of silver for the chloride ion (Sanders et al., 1991). Silver sorbs readily to phytoplankton and to sus- pended sediments. As salinity increases, the degree of
sorption decreases. Nearly 80% of silver sorbed to sus- pended sediments at low salinities desorbs at higher salinities, but desorption does not occur when silver is associated with phytoplankton. Thus, silver incorpora- tion in or on cellular material increases the retention of silver in the estuary, reducing the rate of transport (Sanders & Abbe, 1987).
In California, USA, anthropogenic sources contrib- uted 50% more silver to sediments of coastal basins than did natural sources, as judged by sedimentary basin fluxes of 0.09 µg/cm2 in anthropogenic sources of silver and 0.06 µg/cm2 in natural sources (Bruland et al., 1974). Silver is tightly bound by sewage sludge, and elevated silver concentrations in sediments are often characteristic of areas near sewage outfalls. In the absence of sewage, silver in oxidized sediments is asso- ciated with oxides of iron and with humic substances (Bryan & Langston, 1992).
Riverine transport of silver to the ocean is consider- able; suspended materials in the Susquehanna River, Pennsylvania, USA, containing as much as 25 mg sil- ver/kg, result in an estimated transport of 4.5 tonnes of silver to the ocean each year (US EPA, 1980).
Sensitive marine algae accumulated silver from water containing as little as 2 µg silver/litre (as silver nitrate) to whole-cell burdens as high as 58 mg silver/kg dry weight (Sanders & Abbe, 1987). Dissolved silver speciation and bioavailability were important in deter- mining uptake and retention by aquatic plants (Connell et al., 1991). Silver availability was controlled by the concentrations of free silver ion and of other silver complexes, such as silver chloride (Sanders & Abbe, 1989). Silver uptake by phytoplankton was rapid, proportional to silver concentration, and inversely proportional to water salinity. Silver incorporated by phytoplankton was not lost as the salinity increased, and silver associated with cellular material was largely retained in the estuary (Sanders & Abbe, 1989). Diatoms (Thalassiosira sp.), for example, readily accumulated silver from the medium. Once incorporated, silver was tightly bound to the cell membrane, even after the cells were mechanically disrupted (Connell et al., 1991).
Generally, silver accumulation is primarily attributed to the bioavailability of the free silver ion; however, Rein- felder & Chang (1999) demonstrated that in euryhaline marine microalgae (Thalassiosira weissflogii), AgCl(aq) is the principal bioavailable species of inorganic silver.
In short-term (<1 h) experiments on freshwater green algae (Chlamydomonas reinhardtii) at low free silver ion concentrations (8 nmol/litre), silver uptake increases markedly (up to 4 times) as a function of the chloride ion concentration (5 µmol/litre to 4 mmol/litre) (Fortin &
Campbell, 2000).
The ability to accumulate dissolved silver varies widely between species. Some reported bioaccumulation factors (calculated as milligrams of silver per kilogram fresh weight organism divided by milligrams of silver per litre of medium) are 210 in diatoms, 240 in brown algae, 330 in mussels, 2300 in scallops, and 18 700 in oysters (US EPA, 1980). Silver is the most strongly accumulated of all trace metals by marine bivalve molluscs (Luoma, 1994). Studies with 110mAg suggest that the half-time persistence of silver is 27 days in mussels, 44–80 days in clams, and more than 180 days in oysters (Fisher et al., 1994). In oysters and other bivalve molluscs, the major pathway of silver accumu- lation was from dissolved silver; uptake was negligible from silver adsorbed onto suspended sediments or algal cells, and oysters eliminated adsorbed silver in the faeces (Abbe & Sanders, 1990; Sanders et al., 1990).
Benthic bivalve molluscs can accumulate silver from certain sediments. Sediment-bound silver was taken up by the Baltic clam (Macoma balthica) at 3.6–6.1 times the concentration in calcite sediments, but less than 0.85 times from manganous, ferrous, and biogenic calcium carbonate sediments (US EPA, 1980). In oysters, silver associated with food was unavailable for incorporation, which may be due to the ability of silver to adsorb rapidly to cell surfaces and to remain tightly bound despite changes in pH or enzymatic activity (Connell et al., 1991). Silver concentrations in American oysters (Crassostrea virginica) held in seawater solu- tions containing 1.0 mg silver/litre for 96 h rose from 6.1 to 14.9 mg/kg fresh weight soft parts; in gills, these values were 5.9 and 33.9 mg/kg, respectively (Thurberg et al., 1974). A similar pattern was evident in common mussels (Mytilus edulis) and quahog clams (Mercenaria mercenaria) (Thurberg et al., 1974). Adult surf clams (Spisula solidissima) immersed for 96 h in seawater containing 10 µg silver/litre had 1.0 mg silver/kg fresh weight soft tissues compared with 0.08 mg/kg in con- trols (Thurberg et al., 1974). Uptake of dissolved silver by oysters was higher at elevated temperatures in the range of 15–25 °C (Abbe & Sanders, 1990). American oysters maintained near a nuclear power plant in Mary- land, USA, that discharged radionuclides on a daily basis into Chesapeake Bay accumulated 110mAg; accumula- tions were higher in summer and autumn than in winter and spring (Rose et al., 1988).
Juvenile Pacific oysters (Crassostrea gigas) exposed for 2 weeks to solutions containing 20 µg sil- ver/litre had high silver accumulations in tissues and a reduced capacity to store glycogen; however, after 30 days of depuration, glycogen storage capacity was restored and 80% of the soluble silver and 27% of the insoluble forms were eliminated, suggesting recovery to a normal physiological state. About 70% of the insoluble silver in Pacific oysters was sequestered as silver sulfide, a stable mineral form that is not degradable, thereby limiting the risk of silver transfer through the food-
chain. Most (69–89%) of the silver accumulated in soft tissues of oysters and clams was sequestered in amoebo- cytes and basement membranes; in scallops and mussels, silver was stored in basement membranes and the peri- cardial gland. In all species of bivalve mollusc, seques- tered silver was in the form of silver sulfide. American oysters excreted about 60% of their accumulated silver in soft tissues within 30 days of transfer to silver-free seawater; soluble forms were preferentially eliminated, and insoluble forms retained (Berthet et al., 1992). Inter- species differences in ability to retain silver among bivalve molluscs are large, even among closely related species of Crassostreid oysters. For example, the half- time persistence of silver was about 149 days in Ameri- can oysters but only 26 days in Pacific oysters (ATSDR, 1990).
Marine gastropods exposed to concentrations as low as 1 µg silver/litre for as long as 24 months showed accumulations as high as 34 mg silver/kg fresh weight soft parts; higher exposure concentrations of 5 and 10 µg silver/litre were associated with whole-body burdens as high as 87 mg silver/kg fresh weight (Nelson et al., 1983).
Among arthropods, grass shrimp (Palaemonetes pugio) rapidly incorporate silver dissolved in brackish water in proportion to its concentration, but not from planktonic or detrital food sources containing elevated silver burdens (Connell et al., 1991). Variations in the ability of decapod crustaceans to accumulate 110mAg from seawater are large, as judged by bioconcentration factors that ranged from 70 to 4000 (Pouvreau &
Amiard, 1974). The reasons for this variability are unknown but may be associated with hepatopancreas morphology (Eisler, 1997). It is generally acknowledged that the hepatopancreas or digestive gland is the major repository of silver in decapods (Greig, 1975; Greig et al., 1977a, 1977b), although silver sequestered there may have less effect than in more sensitive tissues. Aquatic insects concentrate silver in relative proportion to envi- ronmental levels (Nehring, 1976) and more efficiently than most fish species (Diamond et al., 1990). Whole- body bioconcentration factors of silver in three species of aquatic insect (calculated as milligrams total silver per kilogram fresh weight tissue divided by milligrams total silver per litre of medium) ranged from 21 to 240 in water containing 30–65 mg calcium carbonate/litre during exposure of 3–15 days; in bluegill (Lepomis macrochirus), this value was less than 1 after exposure for 28 days (US EPA, 1980).
In experiments on rainbow trout (Oncorhynchus mykiss), the uptake of silver has been shown to be via a sodium ion channel situated on the branchial apical membrane of the gills (Bury & Wood, 1999). Uptake of silver by basolateral membrane vesicles was via a carrier-mediated process, which was ATP-dependent,
reached equilibrium over time, and followed Michaelis- Menten kinetics. A P-type ATPase present in the baso- lateral membrane of the gills can actively transport silver (Bury et al., 1999b).
Forsythe et al. (1996) exposed green algae (Selenas- trum capricornutum) and daphnids (Daphnia magna) to silver nitrate concentrations of 1 and 0.5 µg silver/litre, respectively, and reported bioconcentration factors of 4.8 and 61. No significant accumulation was found in blue- gills exposed to 0.5 µg silver/litre. Coleman & Cearley (1974) found that largemouth bass (Micropterus sal- moides) and bluegills accumulated silver; accumulations increased with increasing concentrations of ionic silver and increasing duration of exposure. Bioconcentration factors of 110mAg for various species of teleosts were as high as 40 after 98 days (Pouvreau & Amiard, 1974).
However, plaice (Pleuronectes platessa; a marine floun- der) and thornback rays (Raja clavata) fed nereid poly- chaete worms labelled with 110mAg retained about 4.2%
of the ingested dose after 3 days (Pentreath, 1977), which suggests that the high silver bioconcentration factors reported by Pouvreau & Amiard (1974) may have been due to loosely bound adsorbed silver. Floun- ders (Pleuronectes sp.) held for 2 months in seawater solutions containing 40 µg silver/litre had elevated silver concentrations in the gut (0.49 mg silver/kg fresh weight) but less than 0.05 mg/kg in all other examined tissues (Pentreath, 1977). Similarly, exposed rays (Raja sp.) contained 1.5 mg silver/kg fresh weight in liver, 0.6 mg/kg in gut, 0.2 mg/kg in heart, and 0.005–
0.18 mg/kg in spleen, kidney, and gill filament (Pen- treath, 1977); liver is usually considered the major repository of silver in teleosts (Garnier et al., 1990).
After a 21-day exposure to 14.5 µg silver/litre, silver levels increased 2- to 20-fold in most tissues of four marine teleosts and two marine elasmobranchs, with the highest concentrations occurring in the livers of teleosts and the gills of elasmo branchs. In sculpins (Oligocottus maculosus), salinity markedly affected silver accumula- tion, with tissue-specific levels approximately 6-fold higher at 18‰ than at 30‰. At lower salinities, a neutral AgClaq complex exists in the water, allowing for
increased bioaccumu lation, whereas at higher salinities, only less bioavailable, negatively charged AgCln1–n
complexes (AgCl2–, AgCl32–, AgCl43–) are present (Webb
& Wood, 2000).
Hogstrand et al. (1996) found that rainbow trout exposed to silver nitrate, silver thiosulfate, and silver chloride accumulated silver in the liver. Highest silver levels were found in the livers of trout exposed to 164 mg/litre as silver thiosulfate. In these fish, the hepatic silver concentration was increased 335 times from the control value. The metallothionein (a protein that sequesters metals, reducing their toxicity in tissues) levels in gills and liver increased with the water silver concentration, and the highest level of metallothionein
was found in liver of fish exposed to silver thiosulfate.
The studies revealed that whereas trout accumulated higher silver levels in the liver from high silver thio- sulfate and silver chloride exposures, these compounds were much less toxic than silver nitrate. The authors suggest that the increased induction of metallothionein might explain the apparent lack of toxicity from the high hepatic silver loads in fish exposed to silver thiosulfate and silver chloride.
At concentrations normally encountered in the environment, food-chain biomagnification of silver in aquatic systems is unlikely (Connell et al., 1991; Ratte, 1999). Silver, as thiosulfate-complexed silver at nominal concentrations of 500 or 5000 µg silver/litre, was con- centrated and magnified over a 10-week period in fresh- water food-chains of algae, daphnids, mussels, and fathead minnows (Pimephales promelas) (Terhaar et al., 1977), although the mechanisms of accumulation in this study were imperfectly understood. Sediments contami- nated with silver sulfide, however, do not seem to pose a major route of entry into the aquatic food web. Aquatic oligochaetes (Lumbriculus variegatus) held on sedi- ments containing 444 mg silver/kg dry weight, as silver sulfide, for 28 days had a low bioconcentration factor of 0.18 (Hirsch, 1998b). Fisher & Wang (1998) showed that trophic transfer of silver in marine herbivores, especially mussels, was dependent on the silver assimi- lation efficiency from ingested food particles, feeding rate, and silver efflux rate. Silver assimilation efficiency is usually less than 30% and lower for sediment than for phytoplankton. Silver assimilation efficiency and distri- bution from ingested phytoplankton particles are modi- fied by gut passage time, extracellular and intracellular digestion rates, and metal desorption at lowered pH. The kinetic model of Fisher & Wang (1998) for mussels predicts that either the solute or the particulate pathway can dominate, depending on the silver partition coeffi- cients for suspended particles and silver assimilation efficiency.
6. ENVIRONMENTAL LEVELS
Silver is comparatively rare in the Earth’s crust — 67th in order of natural abundance of elements. The crustal abundance is an estimated 0.07 mg/kg and is predominantly concentrated in basalt (0.1 mg/kg) and igneous rocks (0.07 mg/kg). Silver concentrations tend to be naturally elevated in crude oil and in water from hot springs and steam wells. Anthropogenic sources associated with the elevated concentrations of silver in non-living materials include smelting, hazardous waste sites, cloud seeding with silver iodide, metal mining, sewage outfalls, and especially the photoprocessing industry. Silver concentrations in biota were greater in
organisms near sewage outfalls, electroplating plants, mine wastes, and silver iodide-seeded areas than in con- specifics from more distant sites (Eisler, 1997). Silver concentrations reported prior to the implementation of ultra-clean metal sampling, which began in the late 1980s, should be treated with caution.
Maximum concentrations of total silver that have been recorded in selected non-biological materials are 36.5 ng/m3 in air near a smelter in Idaho, USA (ATSDR, 1990); 2.0 µg/m3 in atmospheric dust (Freeman, 1979);
0.1 µg/litre in oil well brines (US EPA, 1980); 4.5 µg/li- tre in precipitation from clouds seeded with silver iodide; 6.0 µg/litre in groundwater near a hazardous waste site (ATSDR, 1990); 8.9 µg/litre in seawater from Galveston Bay, USA (Morse et al., 1993); 260 µg/litre in the Genesee River, New York, USA — the recipient of photographic manufacturing wastes; 300 µg/litre in steam wells (US EPA, 1980); 300 µg/litre in treated photoprocessing wastewaters; 31 mg/kg in some Idaho, USA, soils (ATSDR, 1990); 43 mg/litre in water from certain hot springs (US EPA, 1980); 50 mg/kg in granite (Fowler & Nordberg, 1986); as much as 100 mg/kg in crude oils; and 150 mg/kg in some Genesee River, USA, sediments (US EPA, 1980). It should be noted that only a small portion of the total silver in each of these com- partments is biologically available. More recent moni- toring studies revealed that levels in the Genesee River, USA, had decreased to below the detection limit (<10 µg/litre) in water samples and that maximum sediment values were 55 mg/kg (DEC, 1993). Gill et al.
(1997) monitored surface waters and municipal and industrial discharge effluents in Colorado, USA, using ultra-clean sampling protocols. Silver measurements in unfiltered samples spanned more than 4 orders of magni- tude, from a high of 33 µg/litre in industrial effluent to a low of 2 ng/litre. In general, upstream of the industrial discharges, unfiltered silver concentrations ranged from 3 to 20 ng/litre; downstream concentrations ranged from 3 ng/litre to 1 µg/litre.
Silver is usually found in extremely low concen- trations in natural waters because of its low crustal abundance and low mobility in water (US EPA, 1980).
One of the highest silver concentrations recorded in fresh water, 38 µg/litre, occurred in the Colorado River, USA, downstream of an abandoned gold–copper–silver mine, an oil shale extraction plant, a gasoline and coke refinery, and a uranium processing facility (US EPA, 1980).
In general, silver concentrations in surface waters in the USA were lower during 1975–1979 than during 1970–1974 (ATSDR, 1990). About 30–70% of the silver in surface waters may be ascribed to suspended particles (Smith & Carson, 1977), depending on water hardness and salinity. The most recent measurements of silver in rivers, lakes, and estuaries using clean techniques show
levels of about 0.01 µg/litre for pristine, unpolluted areas and 0.01–0.1 µg/litre in urban and industrialized areas (Ratte, 1999).
Silver can remain attached to oceanic sediments for about 100 years under conditions of high pH, high salinity, and high sediment concentrations of iron, manganese oxide, and organics (Wingert-Runge &
Andren, 1994). Estuarine sediments that receive metals, mining wastes, or sewage usually have higher silver concentrations (>0.1 mg/kg dry weight) than do non- contaminated sediments. Sediments in Puget Sound, Washington, USA, were significantly enriched in silver, in part from human activities; concentrations were higher in fine-grained particles (Bloom & Crecilius, 1987). Silver levels in sediments from San Francisco Bay, USA, declined from approximately 1.6 mg/kg (15 nmol/g dry weight) in the late 1970s to 0.2 mg/kg (1.8 nmol/g) in the late 1990s (Hornberger et al., 1999).
Marine annelids and clams accumulate dissolved and sediment-bound forms of silver. Uptake of silver from sediments by marine polychaete annelids decreased in sediments with high concentrations of humic substances or copper but increased in sediments with elevated con- centrations of manganese or iron (Bryan & Langston, 1992).
Maximum concentrations of total silver recorded in field collections of living organisms, in milligrams of silver per kilogram dry weight, were 1.5 in liver of marine mammals (Szefer et al., 1994), 2 in liver and 6 in bone of trout from ecosystems receiving precipitation from silver iodide-seeded clouds (Freeman, 1979), 7 in kidneys and 44 in liver of birds from a metals-
contaminated area (Lande, 1977), 14 in marine algae and macrophytes (Eisler, 1981), 30 in whole annelid worms (Bryan & Hummerstone, 1977), 110 in whole mush- rooms (Falandysz & Danisiewicz, 1995), 133–185 in soft parts of clams and mussels near sewage and mining waste outfalls (Luoma & Phillips, 1988; ATSDR, 1990), and 320 in whole gastropods from South San Francisco Bay (Luoma & Phillips, 1988). Silver concentrations in conspecifics from areas remote from anthropogenic contamination were usually lower by 1 or more orders of magnitude. The accumulation of silver by benthic organ- isms from marine sediment is attributed, in part, to the formation of stable complexes of silver with chlorine, which, in turn, favours the distribution and accumulation of silver (Ratte, 1999).
Silver is a normal trace constituent of many organ- isms (Smith & Carson, 1977). In terrestrial plants, silver concentrations are usually less than 1.0 mg/kg ash weight (equivalent to less than 0.1 mg/kg dry weight) and are higher in trees, shrubs, and other plants near regions of silver mining; seeds, nuts, and fruits usually contain higher silver concentrations than do other plant parts (US EPA, 1980). Silver accumulations in marine
algae (maximum 14.1 mg/kg dry weight) are due mainly to adsorption rather than uptake; bioconcentration factors of 13 000–66 000 are not uncommon (ATSDR, 1990; Ratte, 1999).
Silver concentrations in molluscs vary widely between closely related species and among conspecifics from different areas (Bryan, 1973; Eisler, 1981). The highest silver concentrations in all examined species of molluscs were in the internal organs, especially in the digestive gland and kidneys (Eisler, 1981; Miramand &
Bentley, 1992). Elevated concentrations of silver (5.3 mg/kg dry weight) in shells of limpets from uncon- taminated sites suggest that silver may actively partici- pate in carbonate mineral formation (Navrot et al., 1974), but this needs verification (Eisler, 1997). In general, silver concentrations were elevated in molluscs collected near port cities and in the vicinities of river discharges (Fowler & Oregioni, 1976; Berrow, 1991), electroplating plant outfalls (Eisler et al., 1978; Stephen- son & Leonard, 1994), ocean dump sites (Greig, 1979), and urban point sources, including sewage outfalls (Alexander & Young, 1976; Smith & Carson, 1977;
Martin et al., 1988; Anderlini, 1992; Crecelius, 1993), and from calcareous sediments rather than detrital organic or iron oxide sediments (Luoma & Jenne, 1977).
Season of collection (Fowler & Oregioni, 1976; Sanders et al., 1991) and latitude (Anderlini, 1974) also influence silver accumulations. Seasonal variations in silver con- centrations of Baltic clams were associated with sea- sonal variations in soft tissue weight and frequently reflected the silver content in the sediments (Cain &
Luoma, 1990). Oysters from the Gulf of Mexico vary considerably in whole-body concentrations of silver and other trace metals. Variables that modify silver concen- trations in oyster tissues include age, size, sex, repro- ductive stage, general health, and metabolism of the animal; water temperature, salinity, dissolved oxygen, and turbidity; natural and anthropogenic inputs to the biosphere; and chemical species and interactions with other compounds (Presley et al., 1990). Silver concen- trations in whole American oysters from Chesapeake Bay were reduced in summer, reduced at increasing salinities, and elevated near sites of human activity;
chemical forms of silver taken up by oysters included the free monovalent ion and the uncharged AgCl0 (Sanders et al., 1991; Daskalakis, 1996). Declines in tissue silver concentrations of the California mussel (Mytilus californicus) were significant between 1977 and 1990; body burdens decreased from 10–70 mg/kg dry weight to less than 2 mg/kg dry weight, apparently related to the termination of metal plating facilities in 1974 and decreased mass emission rates by wastewater treatment facilities (Stephenson & Leonard, 1994).
Among arthropods, pyrophosphate granules isolated from barnacles have the capability to bind and effec- tively detoxify silver and other metals under natural
conditions (Pullen & Rainbow, 1991). In a Colorado, USA, alpine lake, silver concentrations in caddisflies and chironomid larvae usually reflected silver concen- trations in sediments; seston, however, showed a high correlation with lakewater silver concentrations from 20 days earlier (Freeman, 1979).
In other studies, silver concentrations in fish mus- cles rarely exceeded 0.2 mg/kg dry weight and usually were less than 0.1 mg/kg fresh weight; livers contained as much as 0.8 mg/kg fresh weight, although values greater than 0.3 mg/kg fresh weight were unusual; and whole fish contained as much as 0.2 mg/kg fresh weight.
Livers of Atlantic cod (Gadus morhua) contained significantly more silver than muscles or ovaries; a similar pattern was evident in other species of marine teleosts (Hellou et al., 1992; Szefer et al., 1993).
Accumulation of silver in offshore populations of teleosts is unusual, even among fish collected near dumping sites impacted by substantial quantities of silver and other metals. For example, of seven species of marine fish from a disposal site in the New York Bight, USA, that were examined for silver content, concentra- tions were highest (0.15 mg/kg fresh weight) in muscle of blue hake (Antimora rostrata; Greig et al., 1976).
Similarly, the elevated silver concentration of 0.8 mg/kg fresh weight in liver of winter flounder (Pleuronectes americanus) was from a specimen from the same general area (Greig & Wenzloff, 1977). The similarity between laboratory experiments with snow crab (Chionoecetes opilio) and American plaice (Hippoglossoides plates- soides) and field data from an estuary receiving signi- ficant inputs of anthropogenic silver suggests that predation is the major transfer route of silver to marine benthic predators (Rouleau et al., 2000).
Silver concentrations were lower in muscles of Antarctic birds (0.01 mg/kg dry weight) than in livers (0.02–0.46 mg/kg dry weight) or faeces (0.18 mg/kg dry weight; Szefer et al., 1993). Silver concentrations in avian tissues, especially in livers, were elevated in the vicinity of metal-contaminated areas and in diving ducks from San Francisco Bay, USA.
The concentration of silver in three species of seal collected in the Antarctic during 1989 was highest in the liver (1.55 mg/kg dry weight) and lowest in muscle (0.01 mg/kg dry weight); intermediate values were found in kidney (0.29 mg/kg dry weight) and stomach contents (0.24 mg/kg dry weight; Szefer et al., 1993). The mean concentration of silver in livers from normal female California sea lions (Zalophus californicus), with normal pups, was 0.5 mg/kg dry weight (Martin et al., 1976).
Silver concentrations in tissues of Antarctic seals were related to, and possibly governed by, concentrations of other metals (Szefer et al., 1993). In muscle, silver was inversely correlated with zinc; in liver, silver was posi- tively correlated with nickel, copper, and zinc; and in
kidney, correlations between silver and zinc and between silver and cadmium were negative (Szefer et al., 1993).
Saeki et al. (2001) analysed samples from three species of pinniped from the North Pacific Ocean. In northern fur seals (Callorhinus ursinus), relatively high concen- trations of silver were found in the liver and body hair, with some 70% of the body burden located in the liver.
Hepatic silver concentrations were significantly corre- lated with age in northern fur seals and Steller sea lions (Eumetopias jubatus). Silver concentrations (milligrams per kilogram wet weight) were 0.04–0.55 for northern fur seals, 0.1–1.04 for Steller sea lions, and 0.03–0.83 for harbour seals (Phoca vitulina). Becker et al. (1995) reported silver levels in Alaskan beluga whale (Del- phinapterus leucas) liver to be 2 orders of magnitude higher than for any other marine mammals, although no adverse effects were reported.
7. EFFECTS ON ORGANISMS IN THE LABORATORY AND FIELD
7.1 Aquatic environment
In solution, ionic silver is extremely toxic to aquatic plants and animals (Nehring, 1976; Nelson et al., 1976;
Calabrese et al., 1977a; Gould & MacInnes, 1977; Smith
& Carson, 1977; US EPA, 1980; Buhl & Hamilton, 1991; Bryan & Langston, 1992), and aqueous concentra- tions of 1–5 µg/litre killed sensitive species of aquatic organisms, including representative species of insects, daphnids, amphipods, trout, flounder, and dace. At nominal water concentrations of 0.5–4.5 µg/litre, accumulations in most species of exposed organisms were high and had adverse effects on growth in algae, clams, oysters, snails, daphnids, amphipods, and trout;
moulting in mayflies; and histopathology in mussels (Eisler, 1997).
The acute toxicity of silver to aquatic species varies drastically by the chemical form and correlates with the availability of free ionic silver (Wood et al., 1994). In natural aquatic systems, ionic silver is rapidly com- plexed and sorbed by dissolved and suspended materials that are usually present. Complexed and sorbed silver species in natural waters are at least 1 order of magni- tude less toxic to aquatic organisms than the free silver ion (Rodgers et al., 1994; Ratte, 1999). Thus, silver nitrate, which is strongly dissociated, is extremely toxic to rainbow trout; the 7-day LC50 value is 9.1 µg/litre.
Silver thiosulfate, silver chloride, and silver sulfide were relatively benign (7-day LC50 values >100 000 µg/litre), presumably due to the abilities of the anions to remove ionic silver from solution (Wood et al., 1994, 1996b;
